#826
Effects of feral free-roaming horses
on semi-arid rangeland ecosystems:
an example from the sagebrush steppe
K. W. DAVIES,1, G. COLLINS,2
1
AND
C. S. BOYD1
Eastern Oregon Agricultural Research Center, United States Department of Agriculture, Agricultural Research Service,
Burns, Oregon 97720 USA
2
Sheldon-Hart Mountain National Wildlife Refuge Complex, United States Department of Interior,
Fish and Wildlife Service, Lakeview, Oregon 97630 USA
Citation: Davies, K. W., G. Collins, and C. S. Boyd. 2014. Effects of feral free-roaming horses on semi-arid rangeland
ecosystems: an example from the sagebrush steppe. Ecosphere 5(10):127. http://dx.doi.org/10.1890/ES14-00171.1
Abstract.
Feral horses (Equus caballus) are viewed as a symbol of freedom and power; however, they are
also a relatively unmanaged, non-native grazer in North America, South America, and Australia.
Information about their influence on vegetation and soil characteristics in semi-arid rangelands has been
limited by confounding effects of cattle (Bos taurus) grazing and a lack of empirical manipulative studies.
We compared vegetation and soil surface characteristics in feral horse grazed areas and ungrazed
exclosures at five sagebrush (Artemisia) steppe sites in northern Nevada. Horse grazed areas had lower
sagebrush density and plant diversity, greater soil penetration resistance, and lower soil aggregate stability
than ungrazed areas. Herbaceous cover and density generally did not differ between grazed and ungrazed
treatments, with the exception of heavily grazed sites in which perennial grass cover was reduced. The
cumulative effect of feral horses on soil characteristics suggests that they may affect the ecological function
of semi-arid rangelands by increasing the risk of soil erosion and potentially decreasing availability of
water for plant growth. The two-fold increase in sagebrush density with horse exclusion suggests that feral
horses may limit sagebrush recruitment and thereby negatively impact Greater Sage-grouse (Centrocercus
urophasianus) and other sagebrush associated wildlife. The effects of feral horses on sagebrush and other
semi-arid ecosystems should be considered when developing conservation plans for these ecosystems and
associated wildlife.
Key words: Artemisia; Equus caballus; exclosures; grazing; herbivory; soil aggregate stability; trampling; wild horses.
Received 4 June 2014; revised 16 August 2014; accepted 18 August 2014; final version received 11 September 2014;
published 22 October 2014. Corresponding Editor: R. R. Parmenter.
Copyright: Ó 2014 Davies et al. This is an open-access article distributed under the terms of the Creative Commons
Attribution License, which permits unrestricted use, distribution, and reproduction in any medium, provided the
original author and source are credited. http://creativecommons.org/licenses/by/3.0/
E-mail: [email protected]
INTRODUCTION
States is contentious because they are viewed as a
symbol of freedom and strength and an icon of
the American West (Beever 2003), but they are
also a non-native, feral grazer. The concern with
feral horses is that unmanaged and poorly
managed non-native grazers can have substantial
impacts on ecosystem integrity as seen with
poorly managed livestock in the western United
States (Fleischner 1994, Jones 2000). The potential
Free-ranging horses (Equus caballus) are a
management and conservation concern in North
America (Turner 1987, Beever 2003, Girard et al.
2013) and other locations around the globe (e.g.,
Rogers 1991, Zalba and Cozzani 2004, Nimmo
and Miller 2007, de Villalobos and Zalba 2010).
Management of free-ranging horses in the United
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DAVIES ET AL.
for feral horse effects may be large in some areas
even with low populations because feral horse
grazing is largely unmanaged and on the
landscape year round; where, in contrast, domestic livestock grazing is much more intensively managed through fencing, rotation grazing,
herding, salt and mineral supplementation, and
water sources (Beever 2003). A large portion of
the western United States has the potential to be
affected by feral horses as there are currently
over 40,000 free-ranging horses and burros
occupying 12.8 million ha (BLM 2014).
It is well established that ungulates affect
ecosystems and can modify plant community
composition and structure (Hobbs 1996, Augustine and McNaughton 1998). The effects of
herbivory depend on the complex interaction of
the characteristics of the ungulate, plant community, soil, climate, and landscape (Hobbs 1996).
Thus, some of the controversy regarding feral
horses is the result of a limited number of studies
investigating their effects, especially in arid and
semi-arid ecosystems such as the sagebrush
(Artemisia L.) ecosystem. Beever et al. (2003)
found that disturbance sensitive variables could
collectively be used to separate feral horse
occupied areas from horse removed areas in
multivariate ordination space, despite indistinguishability of the two types of sites using data
on ‘key’ forage species monitored by the Bureau
of Land Management. Horse removed areas had
greater shrub, native grass, and total plant cover
and species richness than horse occupied sites
(Beever et al. 2008). Penetration resistance of the
soil surface (an indicator of soil compaction) was
less where feral horses had been removed
compared to horse occupied sites (Beever and
Herrick 2006). Though these studies (Beever et al.
2003, 2008, Beever and Herrick 2006) provide
valuable information regarding horse occupied
and horse removed areas, they do have some
limitations. Horse occupied and horse removed
areas occurred in separate mountain ranges,
though they had similar slope, aspect, dominant
shrub, fire history, and grazing pressure by cattle
(Bos taurus; Beever et al. 2003, 2008, Beever and
Herrick 2006). Without rigorous experimental
control there are undoubtedly some unaccounted
for confounding variables and cause and effect
relationships cannot be determined. When treatments are not randomly assigned and before and
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after treatment data in both control and experimental plots are not acquired, it is not possible to
attribute differences to the treatments (Beever
and Brussard 2000). Though grazing pressure by
cattle was estimated to be minimal to none across
their study sites (Beever et al. 2003, 2008, Beever
and Herrick 2006), cattle presence may obscure
some of the horse effects or make it difficult to
separate. Therefore, there is a critical need to
evaluate feral horse effects in areas without
domestic livestock. Though covering large landscapes, all these studies (Beever et al. 2003, 2008,
Beever and Herrick 2006) also used the same
study sites. Thus, there is a need for further
analysis of feral horse effects in sagebrush
rangelands to broaden the scope of research to
other areas occupied by feral horses.
In grasslands in Argentina, feral horses reduced grass seed density and richness in the
seedbank (Loydi et al. 2012) and decreased plant
species richness and grass abundance and biomass production (de Villalobos and Zalba 2010).
Feral horse use decreased biomass production in
marshes (Turner 1987, 1988) and islands (Rogers
1991, Seliskar 2003). In contrast, Fahnestock and
Detling (1999) reported feral horses did not have
consistent effects on plant cover and that abiotic
factors were more important than feral horse use
in determining plant community dynamics in the
Pryor Mountain Wild Horse Range in southern
Montana and northern Wyoming. In addition,
plant species diversity was slightly greater in
horse grazed compared to ungrazed sites in the
Pryor Mountain lowlands (Fahnestock and Detling 1999). The inconsistency in the reported
effects of feral horse grazing suggests that
vegetation response likely varies by site characteristics. Furthermore, feral horse effects in more
mesic habitats may contribute little to our
understanding of feral horse effects in sagebrush
(Beever and Aldridge 2011) and other arid and
semi-arid ecosystems.
The purpose of this study was to develop a
better understanding of feral horse effects in
semi-arid rangeland ecosystems. We compared
mountain big sagebrush (Artemisia tridentata spp.
vaseyana (Rydb.) Beetle) steppe communities
where feral horses were excluded for 5 years
with adjacent continuously grazed areas on the
Sheldon National Wildlife Refuge in northern
Nevada. Cattle were excluded from the refuge
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since the early 1990s. We hypothesized that: (1)
perennial herbaceous and sagebrush cover and
density would be lower in feral horse grazed
compared to ungrazed areas, (2) plant species
richness and diversity would be less in grazed
compared to ungrazed areas, and (3) soil surface
penetration resistance would be higher and soil
aggregate stability less in grazed compared to
ungrazed areas.
MATERIALS
AND
Experimental design and measurements
A randomized complete block design was
used to determine the response of mountain big
sagebrush steppe communities to feral horse use
and exclusion. Five blocks, dispersed within the
80,300 ha study area, were selected for this
experiment. Blocks were separated by up to 20
km and varied by elevation, slope, aspect, and
soil. The different blocks included north, south,
east, and west aspects. Elevation of blocks ranged
from 1856 to 1914 m above sea level. Prior to
treatment herbaceous and shrub cover and
density varied by block, but treatments within a
block were similar (P . 0.10). Each block
consisted of two adjacent 50 3 60 m plots
separated by a 10-m buffer. Each block was
within 0.1 km of a riparian area with a perennial
spring or small creek. Both treatment plots were
the same distance from the riparian area at each
block. Treatments were randomly assigned to
one of the plots at each block and were: (1) feral
horse grazing exclosure (ungrazed) and (2)
unrestricted (free-roaming) feral horse grazing.
Exclosures were metal pipe buck-and-pole fences
constructed in 2008 to prevent feral horse use.
Exclosures were constructed with three poles
with the bottom and top pole placed 43 cm and
107 cm above the ground, respectively. Wildlife
were not excluded from the exclosures and
motion sensor cameras recorded no difference
in visits by native ungulates inside compared to
outside exclosures.
Vegetation cover and density was measured in
mid-June 2012 and 2013, the fourth and fifth year
after exclosures were constructed. Herbaceous
cover and density was measured by species in 60
0.2-m2 quadrats in each treatment plot. The 0.2m2 quadrats were placed at 3-m intervals on four
50-m transects (starting at 3 m and ending at 45
m) in each plot (15 quadrats per transect). The
four 50-m transects were placed parallel to each
other at 5-m intervals. Cover of biological soil
crust, bare ground and litter were also measured
in the 0.2-m2 quadrats. Shrub cover by species
was measured on the four 50-m transects using
the line-intercept method (Canfield 1941). Shrub
density by species was measured by counting
plants rooted inside 2 3 50-m belt transects laid
over the 50-m transects. Sagebrush density was
separated into juvenile and mature sagebrush.
Sagebrush was considered to be mature if
METHODS
Study area
We conducted this study from 2008 to 2013 on
Sheldon National Wildlife Refuge in northern
Nevada. The study area is mountainous with
elevations at the study sites ranging around 1900
m above sea level. Climate is typical of the
northern Great Basin with cool, wet winters and
hot, dry summers. Average annual precipitation
was between 300 and 400 mm at the study sites
(NRCS 1998). Annual precipitation in 2008, 2009,
2010, 2011, 2012, and 2013 was 68%, 88%, 148%,
110%, 70%, and 69% of average at the Winnemucca, NV weather station, respectively (Weather
Underground 2014), approximately 180 km
southeast of Sheldon National Wildlife Refuge.
Feral horse populations at Sheldon National
Wildlife Refuge from 2004–2013 ranged from
700 to 1200 individuals. The study area within
Sheldon National Wildlife Refuge was 80,300 ha
with feral horse population ranging from 387 to
673 horses with an average of 528 horses between
2007 and 2013. The study area is an unfenced area
in southern portion of the refuge (Fig. 1) and
horses can travel into and out of the study area.
All study sites were mountain big sagebrush
steppe. However, two of the study sites were
lacking a sagebrush overstory because they were
in the interior of the 1999 Badger Fire. Common
perennial grasses included bluebunch wheatgrass
(Pseduoroegneria spicata (Pursh) A. Löve), Idaho
fescue (Festuca idahoensis Elmer), Thurber’s needlegrass (Achnatherum thurberianum (Piper) Barkworth), bottlebrush squirreltail (Elymus elymoides
(Raf.) Swezey), and Sandberg bluegrass (Poa
secunda J. Presl). Common perennial forbs included milkvetches (Astragalus L.), fleabanes (Erigeron
L.), biscuitroots (Lomatium Raf.), hawksbeards
(Crepis L.), phloxes (Phlox longifolia Nutt. and P.
hoodii Richardson), and lupines (Lupinus L.).
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Fig. 1. Map of the study area (red perimeter polygon) inside of Sheldon National Wildlife Refuge (brown
perimeter polygon) with the location of the refuge identified on a map of Nevada, USA.
Stability Test Kit (Synergy Resource Solutions,
Belgrade, MT). Penetration resistance of the soil
surface was measured at five locations in each
treatment in each block with a Geotester Pocket
Penetrometer (Novatest, Italy).
reproductive stems were present and juvenile if it
did not have reproductive stems. Density data
was used to calculated plant species richness and
diversity. We calculated plant diversity as the
Shannon Diversity Index (H) (Krebs 1998).
Forage utilization was measured in late September of 2009–2013 following the method described
in Anderson and Curreir (1973) to quantify the
level of horse use at each block. Soil aggregate
stability and penetration resistance of the soil
surface were measured in June 2013. Soil
aggregate stability was measured at five locations in each treatment in each block with a Soil
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Statistical analyses
We used repeated measures analysis of variance (ANOVA) models in PROC MIXED in SAS
v.9.2 (SAS Institute, Cary, NC) with year as the
repeated variable to evaluate the response of
variables that were repeatedly measured with
feral horse grazing and exclusion. Blocks and
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Table 1. Average herbaceous utilization by feral horses
at each site measured in September each year using
the method described in Anderson and Curreir
(1973). Sites grouped into grazing levels based on
average utilization levels: low-moderate, ,40%;
high, .40%.
tween treatment means were considered significant at P 0.05. Means are reported with one
standard error.
RESULTS
Perennial grass cover response to grazing
treatments varied by grazing level (Fig. 2; P ¼
0.047). In areas with high levels of grazing,
perennial grass cover was half of ungrazed
exclosures (P ¼ 0.02). We did not find evidence
that perennial grass cover varied between areas
with low-moderate levels of grazing and ungrazed exclosures (P ¼ 0.682). Perennial grass
density did not varied by treatment (Fig. 3) and
did not vary by grazing level (P ¼ 0.227 and
0.291, respectively). We also did not find any
evidence that density (Fig. 3) or cover (Fig. 4) of
any other herbaceous plant functional groups
varied between grazed and ungrazed treatments
or was influenced by grazing level or year (P .
0.05). Total herbaceous cover, bare ground, and
litter did not vary between grazed and ungrazed
treatment (P ¼ 0.195, 0.304, and 0.979). Cover of
biological soil crust did not differ between
treatments (P ¼ 0.332). Grazing level and year
had no effect on total herbaceous cover, bare
ground, litter, and biological soil crust cover (P .
0.05).
Sagebrush and other shrub cover did not differ
between grazed and ungrazed treatments (Fig. 5)
or among years (P . 0.05). In sagebrush
occupied blocks, density of mature sagebrush
was 2-fold greater in ungrazed exclosures compared to the grazed treatment (Fig. 6; P ¼ 0.005).
Sagebrush juvenile density followed a similar
pattern, but statistically there was no difference
between treatments (Fig. 6; P ¼ 0.124). Other
shrub density did not differ between treatments
(P ¼ 0.768). Grazing level and year had no effect
on any of the shrub cover and density responses
(P . 0.05).
Plant species richness did not differ between
ungrazed exclosures (22.6 6 1.2) and grazed
treatments (22.5 6 1.3) (P ¼ 0.990). Plant species
diversity (H) was 1.2-fold greater in exclosures
(1.8 6 0.2) compared to the grazed treatment (1.5
6 0.2) (P ¼ 0.042). Grazing level and year did not
influence plant species richness or diversity (P .
0.05).
Soil aggregate stability averaged 1.5 greater in
Utilization
Site (block)
Mean (%)
SE (%)
Grazing level
Tenmile
Buckaroo
Corral
Smith
Cottonwood
14
61
53
48
29
5
8
4
10
5
low-moderate
high
high
high
low-moderate
block by treatment interactions were treated as
random variables. We also analyzed level of
grazing influence on treatment response. Level of
utilization was separated into two categories
based on average utilization measured in September of 2009–2013: low-moderate (,40% utilization; n ¼ 2) and high (.40% utilization; n ¼ 3)
(Table 1). Our categorization of utilization level
was based on Holechek et al. (1999) synthesis of
grazing studies. Covariance structures used in
the repeated measure ANOVAs were selected
using the Akaike’s Information Criterion (Littell
et al. 1996). Response variables that were not
repeatedly sampled (soil aggregate stability and
penetration resistance) were analyzed with
ANOVAs in PROC MIXED in SAS v. 9.2 (SAS
Institute, Cary, NC). Data that violated assumptions of ANOVAs were log-transformed prior to
analyses. All figures and text report original data
(i.e., non-transformed). For analyses, herbaceous
cover and density were grouped into five
functional groups: perennial grasses, Sandberg
bluegrass, annual grasses, perennial forbs, and
annual forbs. Sandberg bluegrass was treated as
a separate functional group from the other
perennial grasses because it is much smaller in
stature and matures earlier (Davies 2008). Cheatgrass (Bromus tectorum L.) was the only annual
grass detected at the study sites. Total herbaceous
cover was the sum of all the herbaceous species
cover. Shrub cover and density was separated
into sagebrush and other shrubs. The other shrub
category included all shrubs except for sagebrush. The two blocks that did not contain
sagebrush were excluded from analyses of
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Fig. 2. Perennial grass cover (mean þ SE), excluding Sandberg bluegrass, in feral horse grazed (grazed) and
ungrazed (exclosure) treatments split by average grazing utilization levels: low-moderate, ,40% utilization; high,
.40% utilization. Utilization was determined in September in each year. Data presented from the fourth and fifth
years after exclosures were constructed. Lowercase letters indicate difference between treatments (P 0.05) at
that grazing level.
did not affect treatment response (P ¼ 0.200).
ungrazed exclosures compared to the grazed
treatment (Fig. 7; P ¼ 0.002), but was not affected
by grazing level (P ¼ 0.724). Averaged across
blocks, the grazed treatment required 2.5-fold
more force to penetrate the soil surface compared
to ungrazed exclosures (P , 0.001). Grazing level
DISCUSSION
Our results suggest that feral horse grazing can
affect some aspects of semi-arid ecosystems.
Fig. 3. Density (mean þ SE) of different plant functional groups in feral horse grazed (grazed) and ungrazed
(exclosure) treatments. POSA, Sandberg bluegrass; PG, perennial grasses excluding POSA; AG, annual grasses;
PF, perennial forbs; and AF, annual forbs. Data presented from the fourth and fifth years after exclosures were
constructed. Lowercase letters indicate difference between treatments (P 0.05).
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Fig. 4. Cover (mean þ SE) of different cover groups in feral horse grazed (grazed) and ungrazed (exclosure)
treatments. POSA, Sandberg bluegrass; AG, annual grasses; PF, perennial forbs; AF, annual forbs; Therb, total
herbaceous; Bare, bare ground; Litter, ground litter; and BSC, biological soil crusts. Data presented from the
fourth and fifth years after exclosures were constructed. Lowercase letters indicate difference between treatments
(P 0.05).
Similar effects have been reported for other large
herbivores (Hobbs 1996, Augustine and
McNaughton 1998, Olff and Ritchie 1998).
Although previous research (Beever et al. 2003,
2008, Beever and Herrick 2006) suggested that
feral horse use affects the sagebrush ecosystem,
our results are the first study that empirically
measured effects of free-roaming feral horses on
the sagebrush steppe. One limitation of our study
was that our study sites were relatively close to
riparian areas with permanent springs or small
creeks that can concentrate horse use (Crane et al.
Fig. 5. Cover (mean þ SE) of different shrub groups in feral horse grazed (grazed) and ungrazed (exclosure)
treatments. ARTR, big sagebrush; Oshrub, shrubs other than ARTR. Data presented from the fourth and fifth
years after exclosures were constructed. Lowercase letters indicate difference between treatments (P 0.05).
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Fig. 6. Density (mean þ SE) of different shrub groups in feral horse grazed (grazed) and ungrazed (exclosure)
treatments. ARTR(M), mature big sagebrush; ARTR(J), juvenile big sagebrush; Oshrub, shrubs other than ARTR.
Big sagebrush data are only from the three blocks that contained big sagebrush. Data presented from the fourth
and fifth years after exclosures were constructed. Lowercase letters indicate difference between treatments (P 0.05).
1997) and thus, horse effects may dissipate
further from water sources. However, Ganskopp
and Vavra (1986) reported that feral horses
rapidly vacate watering areas after drinking.
Feral horses will also travel long distances (up
to 55 km) from water if forage is scarce and even
disperse up to 8 km from water when forage is
plentiful (Hampson et al. 2010). Herbaceous use
at our study sites spanned a wide range of
utilization levels (Table 1). Thus, our results can
likely be extrapolated to a broad array of
different feral horse use scenarios in arid and
semi-arid ecosystems.
At higher levels of horse use there was about
half the perennial grass cover in grazed compared to ungrazed exclosures; however, herba-
Fig. 7. Soil aggregate stability (left) and soil surface penetration resistance (right) in the feral horse grazed
(grazed) and ungrazed (exclosure) treatments five years after exclosures were constructed. Lowercase letters
indicate difference between treatments (P 0.05).
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ceous cover and density were generally not
different between horse grazed and ungrazed
areas in our study. In contrast, Beever et al. (2008)
reported herbaceous cover and abundance generally varied between horse removed and horse
occupied areas. Contrasting our results of no
difference in perennial grass density, abundance
of perennial grasses was greater in horse excluded areas in grasslands in Argentina (de Villalobos
and Zalba 2010). In our study feral horses may
have been affecting herbaceous cover and density, but recovery of herbaceous plant communities
may not be rapid enough to detect differences
with short-term exclusion of horses. Arid and
semi-arid plant communities can be relatively
slow to recover from disturbances. For example,
sagebrush steppe plant communities can take
several decades once a disturbing agent is
removed for even partial recovery (Sneva et al.
1980, West et al. 1984, Anderson and Inouye
2001). Our study sites had less than half and
about a third the density of perennial grasses and
perennial forbs reported for relatively intact
mountain big sagebrush communities in Davies
and Bates (2010); further suggesting these plant
communities may need more time to recover.
Horses had been removed for two to three times
longer in areas evaluated by Beever et al. (2008),
thus over a longer time period we may see a
similar herbaceous response to feral horse exclusion in our study area. De Villalobos and Zalba
(2010) were also comparing longer term (7 years)
horse exclusion to continuous feral horse grazing. Differences between de Villalobos and
Zalba’s (2010) and our results could also have
been the result of different plant community and
site characteristics.
In contrast to the lack of measured effects of
feral horse grazing exclusion on herbaceous
vegetation abundance, sagebrush density was
two-fold higher in horse exclosures suggesting
sagebrush recruitment may be limited by feral
horses, probably though physical damage of
sagebrush. Consumption of sagebrush may also
be influencing its recruitment as feral horses will
consume small quantities of sagebrush, though
they primarily consume grass (Krysl et al. 1984,
McInnis and Vavra 1987). In contrast to our
results, Beever et al. (2008) did not find a
difference in sagebrush frequency between horse
grazed and horse removed areas. However,
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sagebrush density at our plots was low, about
half of the density Davies and Bates (2010)
reported for relatively intact mountain big
sagebrush communities, suggesting that sagebrush was recovering from past disturbances.
Feral horse effects on sagebrush density may be
more evident when sagebrush is recovering from
a disturbance event; thus, potentially explaining
why we found a difference is sagebrush density
between feral horse grazed and ungrazed areas
and Beever et al. (2008) did not. Beever et al.
(2008) also measured the frequency of sagebrush
using a line-intercept, which may produce a
different result than counting individual plants
in a belt transect as we did.
Non-sprouting shrubs, such as sagebrush,
usually do not recover before herbaceous vegetation after a disturbance (e.g., Davies et al. 2007,
Bates et al. 2011). However, this type of research
has primarily investigated the recovery of plant
communities after fire, not the removal of an
herbivore. Fire often completely removes nonsprouting shrubs, but not herbaceous vegetation
(Davies et al. 2007, Bates et al. 2011). In contrast,
areas where an herbivore was removed may have
a residual population of the non-sprouting shrub
to expedite its recovery. Anderson and Holte
(1981) reported that with the cessation of heavy
grazing by sheep and cattle in areas with
sagebrush that sagebrush cover increased 42%
in the next 15 years while no detectable change
occurred in perennial grass cover, however, 10
years later perennial grass cover had increased
more than 20-fold. Similar to our results, Manier
and Hobbs (2006) found that the long term
exclusion of large herbivores in shrub-steppe
communities in the Rocky Mountains resulted in
increased shrub cover, but did not affect grass
cover. Thus our research and the results from
Anderson and Holte (1981) suggest that sagebrush, if present in the community, may recover
faster than perennial herbaceous vegetation or
alternatively, some herbaceous plant groups may
not change much with herbivore exclusion
(Manier and Hobbs 2006). Additional research
would be valuable to determine the mechanisms
underlying these varying responses to herbivore
removal.
We did not measure a statistically significant
difference in juvenile sagebrush density, which
seems to contradict our results of higher densities
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of mature sagebrush. However, average density
of juvenile sagebrush in ungrazed areas was 7.8fold greater than in horse grazed areas (Fig. 6).
Thus, juvenile sagebrush density results are
similar to mature sagebrush density, but not
statically significantly different due to large
variability. This may be due to the episodic
nature of sagebrush recruitment (Perryman et al.
2001). The lack of a statistically significant
difference may have also been due, in part, to
the way we classified juvenile and mature
sagebrush. Sagebrush was considered mature if
it had reproductive stems, which can occur by
the second growing season (Davies et al. 2014).
Thus, relatively small, young sagebrush were
classified as mature because they had reproductive stems. We saw a similar pattern in sagebrush
cover, where we did not measure a statistically
significant difference in sagebrush cover between
grazed and ungrazed areas, but average sagebrush cover was 1.5-fold greater in grazed
compared to ungrazed areas. Many of the
sagebrush plants were relatively small and thus,
we suspect as sagebrush grows larger that its
cover may increase in ungrazed exclosures
because of the higher density of sagebrush and
become statistically different between ungrazed
and grazed areas. The first year mountain big
sagebrush produces reproductive stems plants
are only about half the height and their canopy
cover is less than a fourth of what they will grow
to in the next couple of years (K. W. Davies,
unpublished data). In support of our speculation,
Beever et al. (2008) found that sagebrush cover
was 1.1 to 2.2 times greater in long term (10–14
yr) horse removed areas compared to horse
occupied areas.
Our results suggest effects of feral horses on
species richness and diversity may vary by plant
community and site characteristics (grazingplant coevolution, climate, soils, etc.). Plant
species diversity was greater in exclosures than
feral horse grazed areas in our study. Beever et al.
(2008) reported that species richness was greater
in feral horse removed compared to horse
occupied areas, but they found no difference in
diversity between areas. Plant species richness
and diversity response to excluding feral horses
has been mixed. Species richness and diversity
were greater in mountain grasslands in Argentina after seven years of feral horse exclusion (de
v www.esajournals.org
Villalobos and Zalba 2010). Other studies have
found that feral horse grazing exclusion promoted a decline in species richness in montane
grasslands in Argentina (Loydi et al. 2012) and
plant diversity in the Pryor Mountain Wild Horse
Range (Fahnestock and Detling 1999). Variation
in intensity and persistence of feral horse grazing
likely also correlates to differences in diversity
responses (de Villalobos and Zalba 2010). Similarly, Olff and Ritchie (1998) in synthesis of the
literature reported that herbivore effects on plant
diversity vary by abundance and type of herbivore, plant community composition, and environment characteristics.
Greater penetration resistance of the soil
surface in feral horse grazed areas suggests that
horse trampling compacted the soil. Beever and
Herrick (2006) reported that penetration resistance of the soil surface was 3–15 times higher in
horse occupied compared to horse removed
areas. This difference was greater than we
observed, but exclosures in our study had only
excluded horses for half to a third of the time that
horses had been absent from horse removed
areas used by Beever and Herrick (2006). The
greater penetration resistance of the soil surface
in feral horse grazed areas than ungrazed
exclosures may be a concern, because the soil
environment influences many ecosystem processes (Beever and Herrick 2006). We found that soil
aggregate stability was 1.5 times greater in feral
horse exclosures compared to grazed plots,
further demonstrating that feral horses were
affecting important soil characteristics. In the
Mojave Desert, removal of feral burros (Equus
asinus L.) and domestic cattle did not produce a
consistent soil aggregate stability response (Beever et al. 2006). However, Beever et al. (2006)
only investigated soil aggregate stability with
one to two years of grazing removal. Soil
aggregate stability may take longer to respond
and this likely explains why we found a
difference with five years of feral horse exclusion.
Horse grazed areas are likely at an elevated
risk of soil erosion compared to horse excluded
areas. Infiltration rates are slower in areas with
higher soil penetration resistance leading to
increased runoff risk (Maestre et al. 2002,
Aksakal et al. 2011). Similarly, a decline in
aggregate stability increases the risk of erosion
(Herrick et al. 2001). Soil erosion has the potential
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October 2014 v Volume 5(10) v Article 127
DAVIES ET AL.
to result in irreversible declines in site productivity and stability and high environmental and
economic costs (Pimentel et al. 1995). This
suggests that the effects of feral horse trampling
may have as much or even more influence on
ecosystems than their selective consumption of
plants. In support of this argument, Turner (1987)
reported in marshes in Georgia that trampling, as
it altered the soils, may be the more destructive
component of grazing by feral horses.
Our results suggest that soil characteristics
may need to recover to facilitate an herbaceous
response. Alternatively, herbaceous vegetation
may not have been substantially affected by feral
horse grazing and a significant change in
herbaceous vegetation with feral horse exclusion
should not be expected. Other authors have also
found that the exclusion of herbivores did not
generally increased herbaceous cover and density (e.g., West et al. 1984, Manier and Hobbs 2006,
Davies et al. 2009). However, Manier and Hobbs
(2006) cautioned that their results should not be
interpreted as evidence that herbivory had not
influenced the herbaceous plant community
because herbivory effects could be long lasting.
Our study sites had less than half the density of
perennial grasses found in relatively intact
mountain big sagebrush communities (Davies
and Bates 2010), thus, feral horses may have
significantly influenced the herbaceous communities at our sites. Soil recovery may have to
occur first because increased soil penetration
resistance and decreased aggregate stability can
reduce the amount of water available for plant
growth (Boxell and Drohan 2009). In addition,
high soil penetration resistance can restrict root
growth (Ehlers et al. 1983, Bengough and Mullin
1991, Villamil et al. 2001). Soil compaction and
trampling by large herbivores can also inhibit
plant diversity (Olff and Ritchie 1998).
Our research agrees with previous research
(Beever et al. 2003, 2008, Beever and Herrick
2006) that suggested feral horse grazing can
cause significant ecological effects in sagebrush
steppe and that these effects on sagebrush may
influence sagebrush associated wildlife species
(Beever and Brussard 2004, Beever and Aldridge
2011). Similarly, feral horses have been reported
to negatively impact native wildlife in Argentina
(Zalba and Cozzani 2004). Our research suggests
that feral horse grazing is limiting the recruitv www.esajournals.org
ment of sagebrush which likely prolongs the time
that the plant community is unsuitable habitat
for sagebrush associated wildlife species. This is
especially concerning because sagebrush habitat
is already limited for many sagebrush associated
wildlife, such as Greater Sage-grouse (Centrocercus urophasianus) (Connelly et al. 2000, Crawford
et al. 2004, Davies et al. 2011) and that fires,
which remove sagebrush, are expected to become
larger and more frequent with global climate
change (Fulé 2008). The role feral horse grazing
plays in limiting sage-grouse and other sagebrush associated wildlife conservation needs to
be considered (Beever and Aldridge 2011).
Additional research investigating sagebrush recruitment and longer term evaluation of the
response of sagebrush and herbaceous cover and
density to feral horse exclusion is needed to fully
understand potential effects of feral horses on
habitat for sagebrush associated wildlife.
Conclusions
Feral horse exclusion increased sagebrush
density and plant species diversity and promoted
recovery of important soil surface characteristics.
Collectively, these results suggests that feral
horse grazing at the utilization levels occurring
in this study can affect the ecological function of
semi-arid rangelands and may degrade the
habitat value of these communities for associated
wildlife. Feral horse grazing elevates the risk of
soil erosion by increasing soil penetration resistance and decreasing soil aggregate stability.
However, our study demonstrated that relatively
short-term feral horse exclusion can initiate
recovery of some variables. Other variables
(e.g., perennial grass density and forb cover and
density) may increase with extended horse
exclusion, but long term studies are needed for
verification. Though feral horse effects likely vary
by intensity and frequency of use as well as a
host of other factors, our results suggest that feral
horses have some ecological impacts on semiarid rangelands across a range of levels of
utilization. Our results agree with Beever and
Aldridge (2011) assertion that feral horses’ value
to society must be weighed against their ecological costs. The collective results from our study
suggest that the effects of feral horses should be
considered when developing conservation plans
for sagebrush steppe rangelands and other semi11
October 2014 v Volume 5(10) v Article 127
DAVIES ET AL.
ecological consequences of feral horse grazing
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Beever, E. A., and J. E. Herrick. 2006. Effects of feral
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Beever, E. A., M. Huso, and D. A. Pyke. 2006.
Multiscale responses of soil stability and invasive
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Beever, E. A., R. J. Taush, and P. F. Brussard. 2003.
Characterizing grazing disturbance in semiarid
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Beever, E. A., R. J. Taush, and W. E. Thogmartin. 2008.
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ACKNOWLEDGMENTS
The authors thank the staff at the Sheldon-Hart
Mountain National Wildlife Refuge Complex for
allowing this research to be conducted on lands they
administer and providing the labor and materials to
build the exclosures. The authors also appreciate Aleta
Nafus, Urban ‘‘Woody’’ Strachan, and all the summer
students for assistance with data collection. We are
also grateful to Erik Beever, Marty Vavra, Rob Sharp
and Dave Ganskopp for reviewing earlier versions of
this manuscript. United States Department of Agriculture is an equal opportunity provider and employer.
Mention of a proprietary product does not constitute a
guarantee or warranty of the product by the USDA or
the authors and does not imply its approval to the
exclusion of other products that may also be suitable.
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